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Forest, Range & Wildland Soils

Acid-base Characteristics of Soils in the Adirondack Mountains, New York

^a E&S Environmental Chemistry, Inc., P.O. Box 609, Corvallis, OR 97339
^b Dep. of Plant Soil and Environmental Sciences, Univ. of Maine, Orono, ME 04469
^c Dep. of Fisheries and Wildlife, Oregon State Univ., Corvallis, OR 97331
^d Dep. of Civil and Environmental Engineering, Syracuse Univ., Syracuse, NY 13244
^e SUNY College of Environmental Science and Forestry, Syracuse, NY 13210
^f NYS Dep. of Environmental Conservation, Bureau of Watershed Management, Albany, NY 12233

^* Corresponding author (tim.sullivan{at}ESenvironmental.com)

	ABSTRACT

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
It is believed that atmospheric deposition of S and N in the Adirondack Mountains of New York has depleted soil-base cation pools, reduced soil base saturation (BS), and contributed to enhanced acidification of soils and surface waters. However, data to determine changes in soil characteristics are generally lacking. It is expected that soil acid-base status will improve as acidic deposition declines in response to atmospheric emissions controls. We studied edaphic characteristics at 199 locations within 44 statistically selected Adirondack lake-watersheds, plus 26 additional watersheds that are included in long-term lakewater monitoring programs. The statistically selected watersheds were chosen to be representative of Adirondack watersheds containing lakes larger than 1 ha and deeper than 1 m that have lakewater acid neutralizing capacity (ANC) less than or equal to 200 µmol_c L^–1. Results of soil analyses were extrapolated to the watersheds of 1320 low ANC lakes. In general, the concentrations of exchangeable base cations, base saturation, and soil pH were low. More than 75% of the target lakes received drainage from watersheds having average B horizon exchangeable Ca concentrations < 0.52 cmol_c kg^–1, base saturation < 10.3%, and pH (H₂O) < 4.5. Variations in the effective cation exchange capacity in both O and B horizons were closely correlated with soil organic matter content. These data provide a baseline against which to compare future changes in regional soil chemistry, and provide input data for aquatic and terrestrial effects models intended to project future changes in surface water chemistry, biological conditions, and forest health.

Abbreviations: AEAP, Adirondack effects assessment program • ALSC, Adirondack Lakes Survey Corporation's survey • ALTM, Adirondack Long-term monitoring project • ANC, acid neutralizing capacity • BS, base saturation • CEC_e, effective cation-exchange capacity • DDRP, Direct Delayed Response Project • ELS, Eastern Lakes Survey • EMAP, Environmental Monitoring and Assessment Program • LOI, loss on ignition • NY-GAP, New York State Gap Analysis Project

	INTRODUCTION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
ECOSYSTEM DAMAGE from air pollution in the Adirondack Mountains, New York, is believed to have been substantial, mainly from atmospheric deposition of S (Sullivan, 2000; Driscoll et al., 2003). Most efforts to quantify damages, and to examine more recent ecosystem recovery, have focused on lakewater chemistry. However, relatively large decreases in regional upwind S emissions and generally similar decreases in S deposition in the Adirondack Mountains over the past two decades have resulted in limited recovery of lakewater acid-base chemistry (Stoddard et al., 1998, 2003; Driscoll et al., 2003). Sulfate concentrations in lakewater have decreased markedly, but so have the concentrations of base cations. Therefore, measured increases in lakewater pH and ANC have generally been small. This limited chemical recovery of surface waters in the Adirondack Mountains and elsewhere in the northeastern USA has been attributed in large part to base cation depletion, especially Ca, of watershed soils (Lawrence et al., 1995; Likens et al., 1996) in response to long-term elevated levels of S deposition. Changes in base cation deposition have also likely contributed (Gbondo-Tugbawa and Driscoll, 2003). The relatively small recent increases in lakewater pH and ANC in response to substantial reductions (>40%) in S deposition (Stoddard et al., 2003) might be attributable to remaining base cation exchange buffering in these watershed soils. There is evidence that historic increases in S deposition resulted in large increases in base cation concentrations of many Adirondack lakes, resulting in only small reductions in ANC during that period (Sullivan et al., 1990; Cumming et al., 1992). Thus, it might be reasonable to expect that subsequent decreases in solution SO₄^2–concentrations and mobility would result in decreased base cation concentrations in runoff. Unfortunately, data that characterize the acid-base chemistry of Adirondack soils are scarce, particularly data that define changes in soil conditions over time.

Databases developed by six major research programs offer an opportunity to evaluate the chronic effects of acidic deposition on acid-sensitive Adirondack watersheds: the Eastern Lakes Survey (ELS; Linthurst et al., 1986), Direct Delayed Response Project (DDRP) (Church et al., 1989), Environmental Monitoring and Assessment Program (EMAP; Larsen et al., 1994), Adirondack Long-Term Monitoring Project (ALTM; Driscoll et al., 2003), Adirondack Effects Assessment Program (AEAP), and the Adirondack Lakes Survey Corporation's (ALSC) survey (Kretser et al., 1989). The ELS, DDRP, and EMAP studies were all statistically based, thereby allowing population estimates to be developed. The ALTM and AEAP involve ongoing long-term lake monitoring efforts, but are not statistically based. Only the DDRP involved sampling and analysis of soil. Today, EMAP provides the best available statistical foundation for Adirondack lake assessment because it included lakes as small as 1 ha in area and the statistical framework was based on more accurate maps than DDRP and ELS. The lack of soil data for EMAP watersheds limits the utility of this robust statistical framework for our purposes.

Before this study, there were no soil chemistry data available for the EMAP sample watersheds in the Adirondack Mountains. Furthermore, there were no soil databases specific to the Adirondack Mountains that were well-suited for evaluating the effects of acidic deposition on soils, or for providing the basis for model projections of aquatic and terrestrial resource recovery. This critical data deficiency regarding Adirondack soils also precluded effective regional modeling to (a) predict future responses of Adirondack watersheds to varying emissions control scenarios, (b) determine critical loads of atmospheric deposition required to protect against further acidification, or (c) define atmospheric deposition goals to allow resource recovery.

The U.S. Environmental Protection Agency's (EPA) DDRP database (Church et al., 1989) has been used for several past modeling studies to project the response of Adirondack lakes to varying scenarios of future acidic deposition (e.g., NAPAP, 1991; Turner et al., 1992; Van Sickle and Church, 1995). However, there are several limitations of the DDRP soils database for conducting an assessment of Adirondack lake responses to various atmospheric deposition scenarios today because:

1. Soil aggregation for the DDRP was based on mapped soils information for the entire Northeast, rather than the Adirondack subregion specifically (a consequence of the broader regional objectives of the DDRP study). As a result, soil data for a particular Adirondack watershed were not necessarily derived from soil pedons excavated within that watershed, or even within the Adirondack region.
   
2. The DDRP soil data are about two decades old and are likely not compatible with lake data collected for current assessment purposes, including data for intensively studied watersheds.
   
3. Although the DDRP watersheds were statistically selected, they are only representative of watersheds containing lakes larger than 4 ha (a consequence of the regional maps available for Eastern Lakes Survey lake selection in 1985). It has subsequently been shown that a high proportion of smaller Adirondack lakes (1–4 ha) are acidic or highly acid-sensitive (Kretser et al., 1989; Baker et al., 1990; Sullivan, 1990; Sullivan et al., 1990), but were not included in the DDRP design.
   

The degree to which regional Adirondack soils exert controls on drainage water chemistry is not well understood, but may be substantial. For example, recent research in Shenandoah National Park, Virginia, has shown a consistent pattern of lower streamwater ANC in watersheds (n = 14) having lower soil BS. All Shenandoah watersheds that had BS < 15% also had average streamwater ANC < 100 µmol_c L^–1. None of the watersheds that had ANC < 20 µmol_c L^–1 showed BS greater than about 12% (Sullivan et al., 2003).

The purpose of our study was to characterize soils from the population of Adirondack watersheds that are potentially sensitive to ecological damage from acidic deposition. The research was undertaken as part of a larger effort to model ecosystem recovery responses to declines in atmospheric S and N deposition. This effort involved soil sampling in a statistically representative group of EMAP watersheds, and in a group of ALTM and AEAP watersheds. Results of this regional soil study are reported here.

	MATERIALS AND METHODS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Site Description
The Adirondack Mountains region, located in northern New York, is a large (2.4 x 10⁶ ha) forested area largely underlain by granitic gneiss and metasedimentary rocks. Spodosols developed from glacial surficial materials cover much of the region. Precipitation is high, averaging 100 cm yr^–1, with greatest values at high elevations and in western sections of the region (Driscoll et al., 1991). Lakes and streams are plentiful. Because of geological and edaphic watershed characteristics combined with a century of acidic deposition, there are about 188 lakes in the EMAP frame that have recently been acidic (having ANC 0 µmol_c L^–1) and an additional 583 that had low summer ANC (between 0 and 50 µmol_c L^–1). In many areas of the Adirondack Park, wet deposition of S is between 6 and 10 kg S ha^–1 yr^–1 and wet deposition of NO₃–N is between 3 and 4 kg N ha ^–1 yr^–1 (Ito et al., 2002). Both upwind emissions and local deposition of S and N have declined in recent years, and are expected to continue to decline in response to implementation of the Clean Air Act Amendments of 1990 and various controls on mobile sources of air pollution (Driscoll et al., 2003; USEPA, 2004).

Watershed and Soil Sampling Site Selection
A group of watersheds was selected for soil sampling based on the EMAP statistical design (Fig. 1). An additional group of watersheds was selected for sampling from among those that are subjects of long-term chemical and biological monitoring efforts in AEAP and ALTM. In the Adirondacks, the regional EMAP probability sample consisted of 115 lakes and their watersheds, representative of 1829 lakes > 1 ha as depicted on 1:100 000 scale topographic maps. Details of the EMAP design were given by Larsen et al. (1994). Whittier et al. (2002) presented an overall assessment of the relative effects of various environmental stressors across northeastern lakes using EMAP probability survey data. Based on EMAP data from 1991–1994, an estimated 42% of the lakes in the Adirondacks had summer index ANC 50 µmol_c L^–1 and another 30% had summer ANC between 50 and 200 µmol_c L^–1. We focused this study on watersheds containing these two strata of low ANC lakes (50 and 50–200 µmol_c L^–1) as they are considered to be most responsive to changes in air pollution. Lakewater ANC is a response variable that integrates watershed acid-base chemistry reflecting biotic, edaphic, geologic, and hydrologic conditions throughout the watershed. We used a random selection process with a systematic spatial component to choose candidate watersheds for soil sampling from among the 44 EMAP sample lake watersheds with ANC 50 µmol_c L^–1 and the 39 EMAP lake watersheds with ANC between 50 and 200 µmol_c L^–1. Both primary and alternate sampling candidate watersheds were selected in the order they were to be included, in anticipation of the problem that some of the selected watersheds would currently be unsampleable (e.g., access difficulty or permission denied). The goal was to sample as least 30 EMAP watersheds containing lakes having ANC 50 µmol_c L^–1 and 10 EMAP watersheds containing lakes having ANC between 50 and 200 µmol_c L^–1. To obtain a spatially balanced subsample, we used county as a spatial clustering variable in a manner identical to that used in the original EMAP probability design (Larsen et al., 1994). For lakes with ANC between 50 and 200 µmol_c L^–1, we used a variable probability factor based on lake ANC class (50–100, 100–150, and 150–200 µmol_c L^–1) to obtain more samples in the lower ANC ranges. No variable probability factors were used for the ANC 50 µmol_c L^–1 lakes. Results of data or model projections for the selected EMAP lake watersheds containing lakes with ANC 50 or 200 µmol_c L^–1 can be easily extrapolated to the entire population of lakes, using the original EMAP sample weights adjusted for this random subsampling procedure.

View larger version (106K): [in this window] [in a new window]   Fig. 1. Locations of soil pits in EMAP watersheds sampled for this study, coded by measured B horizon base saturation.

The forest cover types spatial database used in site selection was derived from the New York State Gap Analysis Project (NY-GAP), which was completed in 1998 based on Landsat-5 thematic mapper satellite imagery data acquired between 1991 and 1993. Thirty-one land cover types were identified by NY-GAP using the National Vegetation Classification System (Grossman et al., 1998; Anderson et al., 1998). These were classified for this project into coniferous, mixed, and deciduous forest types.

In most cases, three target site locations were selected for soil sampling per watershed, with each site being randomly located within the identified predominant landscape units located for practical reasons within 200 m of the lakeshore or inlet stream(s) in the lower 60% of each watershed. Some larger EMAP watersheds contained other study watersheds nested within them. Therefore, these larger watersheds contained more than three target soil-sampling sites. Within each of the three most dominant mapped landscape units in each watershed, we identified a target sampling position using random numbering, a map of hydrography, and the identified forest and soil classes.

Field Sampling
Soils were sampled during the summer of 2003. At each soil-sampling site, an initial evaluation was made to determine whether the selected site was representative of the surrounding soil/vegetation unit, and whether the unit matched the mapped soil drainage and forest type classification. An alternate location was chosen if the initial preselected location was heavily disturbed, located on or very close to a road, under water, or not representative of the surrounding topography. If the initial site was unacceptable, the location of the sampling site was moved 50 m parallel to the stream drainage and a new site was evaluated. If the site failed to correspond with the mapped vegetation type, it was only moved if the targeted site conditions were available close by; otherwise, the landscape conditions were accepted as they occurred in the field and recorded, irrespective of the mapped information. The site was then qualitatively evaluated to determine if it was representative of the sampling unit in terms of canopy cover, slope, and vegetation. If it was not, then a more representative site was selected in the immediate vicinity. Site locations were documented using a global positioning system.

For each pedon, we sampled the O and uppermost 10 cm of the mineral B horizon. Where an E horizon was present, it was typically thin and was assumed to be less reactive than the O and B horizons and was not included in the sampling. With limited resources, we chose to focus on the O and B horizons for our chemical analyses. This focus provided the data needed for future modeling efforts (Cosby et al., 1985; Gbondo-Tugbawa and Driscoll, 2002). Because of the relatively abrupt horizon boundaries in the upper solum of these Spodosols, and the likelihood of acidification effects from air pollution occurring from the top down, we were able to spatially anchor and standardize the mineral soil sampling increment among pedons to a constant and morphologically consistent profile segment. Thus, we sampled all mineral soils as B horizons only, and consistently sampled the upper 10-cm increment of this horizon. In no instance was soil sampled across major horizon boundaries. In a few instances, mineral soil was not present at the selected site.

At each sampling location, we gently brushed off the fresh surface litter without disturbing the O horizon. The entire O horizon, from the surface to the underlying mineral soil boundary, was quantitatively excavated from beneath a 15 x 15 cm template. For sample size considerations, where the O horizon was over 20 cm deep, it was divided in half vertically for sampling; where it was shallow, multiple units were sampled and combined. The E horizon, if present, was gently scraped away and the upper 10 cm of the B horizon was collected. The B horizon was not excavated quantitatively from within the frame, but was sampled using hand digging implements to remove a representative sample from the face of the pedon proportionally across the uppermost 10-cm depth increment of the B horizon.

Samples were labeled according to watershed number (1–100), soil pedon number (1–3), horizon, date of sampling, and name of sampler. Replicate soil samples were identified as such. Samples were placed in plastic bags in the field and transported in coolers back to the field laboratory, where the air-drying process was begun before shipment to the analytical laboratory. Samples of the O horizon and the mineral soil were shipped in coolers to the analytical laboratory at the University of Maine, Orono for processing and analyses.

Laboratory Analysis
Upon arrival at the laboratory, soil samples were unpacked and continued to be air-dried. Air-dried samples were then sieved (6 mm for O horizon, 2 mm for mineral soil) and homogenized by riffle-splitter before being subsampled for analysis. The percentage of air-dry moisture content was calculated to allow for the expression of all data on an oven-dried basis. Soils were oven-dried in a forced draft oven at 70°C for O horizons and 105°C for B horizons. Soil chemical analyses utilized methods widely accepted as appropriate for surface chemical characterization of forest soils. Soil pH was measured in CaCl₂ (0.01 M) and deionized water according to Hendershot et al. (1993) using 2 g of organic or 5 g of mineral soil material in 10 mL of solution. Exchangeable base cations and Al were extracted with 1 M NH₄Cl at a ratio of 2 g of organic soil or 5 g of mineral soil to 100 mL extraction solution and shaken for 1 h (Blume et al., 1990; Fernandez et al., 2003). All extracts were vacuum filtered through Whatman 42 filter paper (Whatman Inc., Clifton, NJ) and analyzed by flame emission (K and Na) or plasma emission spectroscopy (Ca, Mg, and exchangeable Al). Exchangeable acidity was determined by extraction with 1 M KCl (Blume et al., 1990) followed by titration to the phenolphthalein endpoint. Total C and N were determined by LECO combustion. As described in Fernandez et al. (2003), soil organic matter content was estimated on oven-dried samples by loss-on-ignition (LOI) at 450°C. Effective cation-exchange capacity (CEC_e) was calculated as the sum of the exchangeable base cations (Ca, Mg, K, and Na) plus exchangeable acidity. Base saturation was calculated as the percentage of the CEC_e occupied by exchangeable bases, whereas Ca, Mg, or Al saturation was calculated as the percentage of the CEC_e occupied by these individual exchangeable cations. The methods used in this study were similar to those used previously in this laboratory defined in the QA Plans and methods of the USEPA Watershed Manipulation Project (Erickson et al., 1991), USDA Forest Service Forest Response Project (Robarge and Fernandez, 1986), and the DDRP of the USEPA (Blume et al., 1990).

Spatial Aggregation
Landscape units were defined on the basis of the 18 soil classes and 3 forest classes identified for this study. Most study watersheds contained several of these landscape units, of which the most dominant (typically n = 3) were sampled for soil chemistry. Results of chemical analyses were spatially aggregated to yield estimates of average conditions for each watershed. These estimates were areally weighted, using data collected from the pedons excavated within the sampled dominant landscape units in the subject watershed plus nearest-neighbor data for the landscape units that were not sampled in the subject watershed.

Quality Assurance
Quality assurance samples included field replicates, sample replicates, blanks, repeated analyses of an in-house standard that was a field composite sample, certified reference materials for instrumental precision, and spiked recovery as appropriate. During the first several days of sampling, composite soil samples were prepared for both O and mineral B horizons. These composites consisted of air-dried, homogenized, and sieved subsamples from eight soil pedons excavated in four different watersheds. After air-drying and thoroughly homogenizing each of the two composites, each was partitioned into multiple sealed plastic bags and kept refrigerated until shipped to the analytical laboratory. Each week during the sampling period, a shipment of soil samples was packaged in coolers and one of each of the composite samples (O and B horizon) was added to the shipment. Results of repeated analyses of each of these composite soil samples (Table 2) provide important information regarding analytical reproducibility throughout the study. The coefficient of variation for most analytes was less than about 10%. Exceptions to this pattern included mainly analytes having very low concentration (e.g., exchangeable K and Na).

View larger version (17K): [in this window] [in a new window]   Fig. 2. Comparison of base saturation results between original (A sample) and replicate (B sample) soil samples. O horizon data are depicted as closed circles and B horizon data as open circles.

	RESULTS AND DISCUSSION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

Eight general soil classes were sampled. The most common were well-drained ablation and basal till (26%) and excessively drained shallow soil over bedrock (24%). None of the other six soil classes accounted for more than 13% of the total (Table 3).

The overall mean soil exchange phase composition from this sample population of O and B horizons for all watersheds sampled is shown in Fig. 3. These data provide a general representation of the contrasts between O and underlying mineral B horizons that are characteristic of cool, temperate climate forest soils. The O-horizons were highly acidic, being dominated by weak organic acidity despite the relatively high BS (45.2%). All of the O horizon exchangeable base cation concentrations were higher than mineral soil values. Exchangeable Ca dominated in these O horizons and the mean exchangeable Ca/Al ratio for these data in the O horizon was 2.3 to 1. Exchangeable acidity in the O horizon, operationally defined by the methods employed, was dominated by exchangeable H⁺, reflecting the importance of organic functional groups in these highly organic soil materials. Mean B horizon CEC_e was 8.3 cmol_c kg^–1 with a BS of 10.0%, compared with 29.6 cmol_c kg^–1 in the O horizon, with a BS of 45.2%. While there was a strong contribution of organic acidity in the B horizon, as evidenced by the considerable amount of exchangeable H⁺, exchangeable Al clearly dominated the cation exchange complex in the mineral soil and the mean B horizon exchangeable Ca/Al ratio was only 0.13.

View larger version (36K): [in this window] [in a new window]   Fig. 3. Mean composition of the effective cation-exchange capacity (CECe) and pHs from all watershed samples analyzed in the Adirondack watershed sampling conducted in 2003.

View larger version (20K): [in this window] [in a new window]   Fig. 4. Relationship between watershed-aggregated B horizon soil base saturation and lakewater ANC for the 70 study lake watersheds.

In a study of episodic stream acidification processes in eight small watersheds (11–967 ha) throughout the northeastern USA, Lawrence (2002) found that exchangeable Ca concentrations in the O_a horizon were strongly correlated with high-flow stream ANC (r² = 0.84). In contrast, we found no significant relationship between average lakewater ANC and exchangeable Ca in the O horizon (p = 0.25).

Soil Acid-Base Chemistry
Acid-sensitive soils, defined here as having low BS (<10%) in the B horizon, were found to be widely distributed throughout the western Adirondacks. The common occurrence of pedons having low B horizon BS (Fig. 1) is important, in part, because concentrations of inorganic Al in soil solution increase as BS decreases, with the increase in dissolved Al being pronounced when the BS falls below 15% (Reuss, 1983; Cronan and Schofield, 1990).

Population Statistics
The distributions of soil acid-base chemical characteristics within Adirondack lake watersheds containing potentially acid-sensitive lakes are given in Table 4. These distributions were developed from the aggregated data for each of the statistically selected study watersheds. The total number of target Adirondack lakes included in the EMAP frame was 1829 (SE = 244). These include the lakes depicted on 1:100 000-scale USGS maps that were larger than 1 ha, deeper than 1 m, and that contained more than 1000 m² of open water. Of those target lakes, an estimated 509 had lakewater ANC > 200 µmol_c L^–1; these were considered insensitive to acidic deposition effects and were not included in the study reported here. The remaining 1320 (SE = 102) low-ANC lakes constituted the frame for extrapolation of soil results. An estimated 58% of these low-ANC target lakes had ANC < 50 µmol_c L^–1 at the time of EMAP sampling (1991–1994).

Acidification Processes
Soil base cation depletion can occur if the leaching of base cations from the soil by acidic deposition occurs faster than the combined replenishment by rock weathering and base cation deposition from the atmosphere. Long-term depletion of soil base cations, reflected in the low soil BS values, represents a natural pattern of soil formation in humid environments. Modern atmospheric chemical inputs of S and N have accelerated the rate of cation leaching, and therefore base cation loss, in watersheds where base cation acid neutralization mechanisms have not been depleted. After exposure to long-term elevated deposition of atmospheric pollutants, soil base cation acid neutralization mechanisms become less effective as BS declines and acid neutralization through Al mobilization becomes increasingly important, as does exchangeable Al on the CEC_e (Norton et al., 2004; Fernandez et al., 2003). Soils in the population studied here from the Adirondack region appear to typify these characteristics, with low pH and BS, and relatively high exchangeable Al. This presumed accelerated soil acidification may have resulted in a fundamental alteration of soil processes, potentially impacting the nutritional status of vegetation and the extent to which soils and drainage waters can recover from acidification. Indirect chemical effects of soil base cation depletion include the acidification of soil solution and increased mobilization of potentially toxic inorganic monomeric Al from soil to solution (Cronan and Grigal, 1995). Such effects have been linked with the observed decline of red spruce (Picea rubens) at high elevations throughout the northeastern USA (DeHayes et al., 1999) and of sugar maple (Acer saccharum) in Pennsylvania (Horseley et al., 2000).

The focus of this study has been on the implications of BS and the composition of the CEC_e in O and B horizons of Adirondack forested watersheds for soil and surface water acid-base status. Two parameters in this data set provide information on the amount of organic matter in these soils. Soil total C is primarily found in organic matter, although some calcareous parent materials undoubtedly contribute to total C values, particularly at high pH. Also, loss on ignition (LOI) is considered an estimate of soil organic matter, although it is taken to be less accurate than direct C measurements. In these soils, LOI proves to be a good estimator of total soil C and by inference, organic matter content (LOI = 1.5 x TC + 2.44; r² = 0.99; where both LOI and TC are given in percentages).

The relationship between total C and CEC_e in both O and B horizons is shown in Fig. 5. Given that the Adirondack region is dominated by coarse textured parent materials of glacial origin, it is evident that variations in soil CEC_e have less to do with clay content than with soil organic matter content in these soil materials. Similar results were found in the Catskill Mountains of New York by Johnson et al. (2000). The correlations found here are relatively strong considering that these data represent total soil C, with some minor contribution of carbonate minerals, and without any consideration of differences in the degree of organic matter humification that could govern the degree of surface reactivity, and thus influence CEC_e. Organic matter derived CEC_e in these soils takes on additional importance when we consider that essentially all of the CEC_e on soil organic materials is pH-dependent. We can expect that there will be, and perhaps have been, significant changes in CEC_e to the extent that soils have been influenced by acidic deposition (Lapenis et al., 2004). In addition, the quality of organic matter and the degree of humification are governed by plant community composition, productivity, and rate of microbial processing in soil. All of these factors are expected to be influenced by a changing physical climate.

View larger version (31K): [in this window] [in a new window]   Fig. 5. Correlations between total C and effective cation exchange capacity (CECe) from all watershed samples analyzed in the Adirondack watershed sampling conducted in 2003.

View larger version (28K): [in this window] [in a new window]   Fig. 6. Correlations between total C in the soil and H+ in equilibrium soil solutions from the soil pH in water measurements.

Comparison with Direct Delayed Response Project Data from the 1980s
To compare distributions of watershed soil chemistry between our study and those of the 1980s DDRP project, we constructed subsets of both datasets so they defined a common population of lakes. The original DDRP population had been constrained by excluding lakes with area <4 ha, maximum depth <1.5 m, watershed area >30 km², and ANC > 400 µmol_c L^–1 (Church et al., 1989). Because our soil survey of EMAP lake watersheds only included lakes with ANC 200 µmol_c L^–1, for this comparison we excluded DDRP lakes with ANC > 200 µmol_c L^–1 and excluded lakes from our EMAP sample that exceeded DDRP lake area, depth, and watershed area limits. Using these data subsets (n = 32 for EMAP; n = 36 for DDRP) and the sample weights from each survey, cumulative distribution functions were constructed so that we could compare the EMAP and DDRP population distributions of lake watersheds for lakewater ANC, and B horizon soil BS and exchangeable Ca (Fig. 7). Differences between the two surveys in terms of population estimates (percentage of lake watersheds) for specific criteria values were evaluated for significance using a z-test (z = difference between estimates/pooled standard error). The pooled standard error was calculated as the square root of the sum of the squared standard error of the estimate for each survey.

View larger version (28K): [in this window] [in a new window]   Fig. 7. Cumulative distribution functions for a common population of Adirondack lake watersheds constructed for DDRP data collected in the mid 1980s and data compiled or collected in this study for EMAP lake watersheds in the early 1990s (lake chemistry) and in 2003 (soil chemistry). Comparisons are shown for (A) lakewater ANC, (B) B horizon soil base saturation, and (C) B horizon soil exchangeable Ca2+.

These populations of Adirondack lake watersheds defined by the DDRP study in the mid 1980s and our analyses of EMAP lake watersheds in the 1990s (water chemistry) and 2003 (soil chemistry) are not totally comparable. For example, our soil analyses were standardized to the top 10 cm of B horizon soil whereas DDRP upper mineral soil samples exhibited varying depths (Church et al., 1989). Nevertheless, the data shown in Fig. 7 suggest that while lakewater chemistry was improving subsequent to large decreases in acidic deposition (Fig. 7A, a result also shown in other studies, Driscoll et al., 2003), Adirondack soil acid-base chemical conditions may have been continuing to deteriorate in most of the acid-sensitive watersheds (Fig. 7 B and 7C). Such an effect would be expected to restrict the extent to which surface water chemistry will be able to recover from acidification in the future and may contribute to future adverse effects on forest soils and vegetation.

	CONCLUSIONS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
A study of Adirondack soil chemistry was undertaken to provide the basis for regional soil characterization, assessment of soil/surface water linkages, and watershed modeling. This study not only provides information regarding regional soil conditions for low-ANC (<200 µmol_c L^–1) Adirondack lake-watersheds (including those potentially susceptible to adverse impacts from acidic deposition), but it also provides the necessary soil data with which to calibrate and implement S and N effects models used in state and national policy development. These data provide a regionally representative soil database for the Adirondacks that will serve as a benchmark against which future soil conditions can be compared. The establishment of a high quality and documented watershed soil database for this vulnerable and highly valued region is critical to provide meaningful assessments of recovery, or lack thereof, in the future. The risk of continued environmental degradation in this region appears real since effects models generally suggest declining soil BS in the future unless acidic deposition is dramatically reduced (Gbondo-Tugbawa and Driscoll, 2002).

	ACKNOWLEDGMENTS

 
This research was funded by the New York State Energy Research and Development Authority, through the Environmental Monitoring, Evaluation, and Protection (EMEP) Program. We thank Mark Watson for his advice and encouragement.

Received for publication March 10, 2005.

	REFERENCES

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 

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